IAEA-TECDOC-1660
Exposure of the Public from Large Deposits of Mineral Residues
INTERNATIONAL ATOMIC ENERGY AGENCY VIENNA
ISBN 978–92–0–116410–0 ISSN 1011–4289
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Exposure of the Public
from Large Deposits of Mineral Residues
AFGHANISTAN ALBANIA ALGERIA ANGOLA ARGENTINA ARMENIA AUSTRALIA AUSTRIA AZERBAIJAN BAHRAIN BANGLADESH BELARUS BELGIUM BELIZE BENIN BOLIVIA
BOSNIA AND HERZEGOVINA BOTSWANA
BRAZIL BULGARIA BURKINA FASO BURUNDI CAMBODIA CAMEROON CANADA
CENTRAL AFRICAN REPUBLIC CHAD CHILE CHINA COLOMBIA CONGO COSTA RICA CÔTE D’IVOIRE CROATIA CUBA CYPRUS
CZECH REPUBLIC DEMOCRATIC REPUBLIC OF THE CONGO DENMARK
DOMINICAN REPUBLIC ECUADOR
EGYPT EL SALVADOR ERITREA ESTONIA ETHIOPIA FINLAND FRANCE GABON GEORGIA GERMANY
GHANA GREECE GUATEMALA HAITI HOLY SEE HONDURAS HUNGARY ICELAND INDIA INDONESIA
IRAN, ISLAMIC REPUBLIC OF IRAQ
IRELAND ISRAEL ITALY JAMAICA JAPAN JORDAN KAZAKHSTAN KENYA
KOREA, REPUBLIC OF KUWAIT
KYRGYZSTAN LATVIA LEBANON LESOTHO LIBERIA
LIBYAN ARAB JAMAHIRIYA LIECHTENSTEIN
LITHUANIA LUXEMBOURG MADAGASCAR MALAWI MALAYSIA MALI MALTA
MARSHALL ISLANDS MAURITANIA MAURITIUS MEXICO MONACO MONGOLIA MONTENEGRO MOROCCO MOZAMBIQUE MYANMAR NAMIBIA NEPAL
NETHERLANDS NEW ZEALAND NICARAGUA NIGER NIGERIA
NORWAY OMAN PAKISTAN PALAU PANAMA PARAGUAY PERU PHILIPPINES POLAND PORTUGAL QATAR
REPUBLIC OF MOLDOVA ROMANIA
RUSSIAN FEDERATION SAUDI ARABIA SENEGAL SERBIA SEYCHELLES SIERRA LEONE SINGAPORE SLOVAKIA SLOVENIA SOUTH AFRICA SPAIN
SRI LANKA SUDAN SWEDEN SWITZERLAND
SYRIAN ARAB REPUBLIC TAJIKISTAN
THAILAND
THE FORMER YUGOSLAV REPUBLIC OF MACEDONIA TUNISIA
TURKEY UGANDA UKRAINE
UNITED ARAB EMIRATES UNITED KINGDOM OF GREAT BRITAIN AND NORTHERN IRELAND UNITED REPUBLIC OF TANZANIA
UNITED STATES OF AMERICA URUGUAY
UZBEKISTAN VENEZUELA VIETNAM YEMEN ZAMBIA ZIMBABWE
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IAEA-TECDOC-1660
EXPOSURE OF THE PUBLIC FROM LARGE DEPOSITS
OF MINERAL RESIDUES
INTERNATIONAL ATOMIC ENERGY AGENCY
VIENNA, 2011
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EXPOSURE OF THE PUBLIC FROM LARGE DEPOSITS OF MINERAL RESIDUES IAEA, VIENNA, 2011
IAEA-TECDOC-1660 ISBN 978-92-0-116410-0
ISSN 1011-4289
© IAEA, 2011
Printed by the IAEA in Austria June 2011
FOREWORD
All minerals and raw materials contain radionuclides of natural origin. In most situations, the exposure of humans to such radionuclides is considered to be part of the normal natural radiation background and is not generally of concern. In some cases, however, the radionuclide concentrations are elevated above normal levels or become elevated as a result of mineral processing activities, and measures for protecting against exposure to the material involved may need to be considered. The mineral or raw material is then treated as radioactive material for the purposes of radiation protection and falls within the definition of naturally occurring radioactive material (NORM).
The IAEA has developed criteria for determining which materials need to be considered for regulatory control. For materials containing only radionuclides of natural origin, the criteria are an activity concentration of 1 Bq/g for 238U, 235U, 232Th and their decay progeny and an activity concentration of 10 Bq/g for 40K. These values were determined on the basis of the activity concentrations of these radionuclides in normal rocks and soil, and represent the (rounded) upper bounds of the ranges of such concentrations as determined by the United Nations Scientific Committee on the Effects of Atomic Radiation (UNSCEAR). The values are intended to apply to all solid materials except foodstuffs, material in transport and radioactive residues in the environment (for which separate criteria apply) and 40K in the body (which is excluded entirely from regulatory requirements).
While the radiation dose was not a consideration in the determination of the above-mentioned regulatory criteria, the IAEA has noted that doses received by individuals as a consequence of the use of these criteria are unlikely to exceed about 1 mSv in a year, excluding the emanation of radon. However, in the case of bulk volumes of material contaminating water pathways, such as large deposits of NORM residues from mining and mineral processing, it has been suggested that case by case evaluation of the dose may be required. It was therefore decided to conduct further investigations of the doses expected to be received as a result of exposure of members of the public to a large NORM residue deposit, with consideration being given to all potentially significant exposure pathways including those involving contamination of water. The investigations were conducted using an evidence based approach involving the review of available information from real world examples of actual NORM residue deposits, as well as a calculation approach involving the modelling of radionuclide migration from a
‘representative’ large NORM residue deposit.
The investigations were carried out during 2009 under contract to the IAEA by a Canadian consulting company, SENES Consultants Limited. This report gives the results of those investigations. The IAEA Officer responsible for the preparation of this report was D.G. Wymer of the Division of Radiation, Transport and Waste Safety.
EDITORIAL NOTE
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CONTENTS
1. INTRODUCTION ... 1
1.1. Background ... 1
1.2. Objective ... 2
1.3. Scope ... 2
1.4. Structure ... 3
2. EMPIRICAL EVIDENCE ... 4
2.1. Types of residue considered ... 4
2.2. Sources of data ... 5
2.3. Characteristics of residues ... 5
2.3.1. Physical and chemical characteristics ... 5
2.3.2. Radioactivity content ... 6
2.3.3. Leachate ... 8
2.3.4. Radionuclide transport in groundwater ... 15
2.3.5. Distribution coefficients ... 16
2.4. Leachate characteristics and distribution coefficients for a representative NORM residue deposit ... 18
2.5. Reported doses arising from NORM residue deposits ... 19
3. DOSE CALCULATION ... 21
3.1. Characteristics of the representative NORM residue deposit ... 21
3.1.1. Physical characteristics ... 21
3.1.2. Distribution coefficients ... 22
3.1.3. Leach rates ... 22
3.1.4. Air emissions ... 22
3.2. Exposure pathways ... 22
3.3. Exposure and uptake ... 24
3.4. Dose ... 24
3.4.1. Groundwater, ingestion of garden and agricultural products, and surface water pathways ... 25
3.4.2. Ingestion and inhalation of dust ... 27
3.4.3. External exposure ... 27
3.4.4. Results ... 29
4. DISCUSSION AND CONCLUSION ... 30
4.1. Conservatism in the dose calculations ... 30
4.2. Comparison with doses determined from measured radionuclide concentrations in water ... 30
4.3. Acid generating NORM residue deposits ... 31
4.4. Disequilibrium in NORM residue deposits ... 31
4.5. Sensitivity of the results to the values of the distribution coefficients ... 31
4.6. Overall conclusions ... 32
APPENDIX I. DOSE CALCULATION ... 33
APPENDIX II. RADIONUCLIDE CONCENTRATIONS IN AIR ... 38
APPENDIX III.SENSITIVITY OF THE DOSE CALCULATIONS TO VARIATIONS IN DISTRIBUTION COEFFICIENTS ... 42 REFERENCES ... 45 CONTRIBUTORS TO DRAFTING AND REVIEW ... 49
1. INTRODUCTION 1.1. Background
All minerals and natural raw materials contain the so-called ‘primordial’ radionuclides 238U,
235U and 232Th and their decay progeny, as well as 40K. In the majority of situations, the radionuclide concentrations are not sufficiently elevated to pose a radiological hazard and the material is not treated as being radioactive for purposes of radiation protection. In some cases however, where the radionuclide concentrations are significantly higher than the normal range of background levels, there may be a potential for exposures that are of concern from a radiation protection point of view. In such cases, the material is treated as radioactive material and thus falls within the definition of NORM.
A wide range of activity concentrations in a wide variety of materials is reported (see, for instance, Ref. [1]). Examples of ores that have been found to be associated with elevated radionuclide concentrations include those of uranium, tin, tantalum, niobium, rare earths, aluminium, copper, gold and phosphate. The mining and processing of these ores can lead to further increases in radionuclide concentrations in the products, by-products or residues. A few examples of the activity concentrations involved are given in Table 1.
TABLE 1. EXAMPLES OF ACTIVITY CONCENTRATIONS IN NORM (adapted from Ref. [2])
Radionculide with highest activity concentration
Typical activity concentration (Bq/g)
Monazite sand 232Th series 40–600
Metal ores, e.g., Nb, Ta, Cu, Au 238U and 232Th series Up to 10
Zircon sand 238U series 2–4
Phosphate rock 238U series 0.03–3
TiO2 feedstocks 232Th 0.001–2
Bauxite 232Th series 0.035–1.4
Red mud (alumina production) 238U, 232Th 0.1–3
Phosphogypsum (H2SO4 process) 226Ra 0.015–3
Niobium extraction slag 232Th 20–120
Tin melting slag 232Th 0.07–15
Scale (oil and gas production) 226Ra 0.1–15 000
Residue (rare earth extraction) 228Ra 20–3 000
Scale (TiO2 pigment production 228Ra, 226Ra <1–1600
Scale (rare earth extraction) 226Ra, 228Th 1000
Sludge (oil and gas production) 226Ra 0.05–800
Residue (niobium extraction) 228Ra 200–500
Coal 238U and 232Th series 0.01–0.025
Scale (coal mines with Ra rich inflow water) 226Ra, 228Ra Up to 200
Activity concentration criteria for the regulation of materials have been established by the IAEA and are given in Ref. [3]. In developing these criteria, it was concluded that it was not practical to derive criteria for radionuclides of natural origin on the basis of dosimetric considerations (as was done for radionuclides of artificial origin), since in many cases such an approach would produce activity concentrations lower than those occurring in the natural environment. Therefore, the activity concentration criteria for radionuclides of natural origin were based on the upper bound of the worldwide distribution of natural radionuclides as given, for example, in Ref. [1]. Consequently, it is stated in Ref. [3] that “It is usually unnecessary to regulate…” material containing radionuclides of natural origin at activity concentrations below 1 Bq/g for radionuclides in the uranium and thorium decay series and below 10 Bq/g for 40K.
1.2. Objective
While dose was not the key consideration in deriving the activity concentration criteria for radionuclides of natural origin, it is stated in Ref. [3] that “Doses to individuals as a consequence of these activity concentrations would be unlikely to exceed about 1 mSv in a year, excluding the contribution from the emanation of radon, which is dealt with separately in the BSS”. It is generally recognized that the main potential for exceeding an annual dose of 1 mSv per unit activity concentration of 1 Bq/g arises from a scenario involving exposure of members of the public to large mineral residue deposits, such as mine tailings and ‘waste’
rock (rock with low levels of mineralization that does not necessarily fall within the IAEA’s definition of waste because of the potential for further use).
The objective of this report is, firstly, to present the findings of an investigation to determine the doses expected to be received by members of the public exposed to large NORM residue deposits, with consideration being given to all potentially significant exposure pathways, including those involving contamination of water, and to the radionuclide activity concentrations in the residue material. The investigation was carried out under contract to the IAEA by SENES Consultants Limited [4] using an evidence based approach involving the review of available information from real world examples of actual NORM residue deposits, as well as a calculation approach involving the modelling of radionuclide migration from a
‘representative’ large NORM residue deposit.
Secondly, the objective of this report is to establish, through the findings of the investigation, the consequences of applying the recommended activity concentration criterion of 1 Bq/g in situations where individuals are exposed to large mine residue deposits. Although, as stated above, the derivation of the activity concentration criterion was not based on dose considerations, a knowledge of the doses likely to be received in such situations would help to establish whether or not the use of the 1 Bq/g criterion for determining the scope of regulatory control could, in any reasonable circumstances, lead to an exposure situation that would be regarded as unacceptable.
1.3. Scope
The scope of the investigation was to assess the doses expected to be received by members of the public living near a large mineral residue deposit using, to the extent possible, an evidence based approach based on data gathered from actual mineral residue deposits. The key assumptions concerning the characteristics of the residue deposit were:
• A nominal deposit volume of 2 million m3 covering 10 ha;
• The presence of radionuclides in the 238U and/or 232Th decay series, each at a concentration of 1 Bq/g;
• The possibility of the residue material being acid generating.
The dose calculations were done using, where appropriate, the same modelling approach and assumptions used for deriving the activity concentration criteria reported in Ref. [3] for radionuclides of artificial origin. The modelling approach and assumptions are described in Ref. [5].
The key elements of the study were:
• To consider the available information from real world examples of actual NORM residue deposits that could be used to define the characteristics of a representative NORM residue deposit and the parameters controlling the migration of radionuclides from the deposit;
• To define the relevant exposure pathways and conduct a dose assessment for the most highly exposed members of the public.
For this assessment, the exposed individuals were assumed to live within a few metres of the NORM residue deposit and, for consistency with the modelling approach used in Ref. [5], an adult and a 1–2 year old child were chosen as hypothetical receptors. The exposure pathways considered in this assessment were:
• Inhalation of airborne dust;
• External exposure from dust deposited on the ground at the residence;
• Ingestion of dust from dust deposited on the ground at the residence;
• Ingestion of garden and agricultural products from irrigation with contaminated groundwater;
• Ingestion of contaminated groundwater;
• Ingestion of fish obtained from contaminated surface water.
For consistency with the approach taken in Refs [3, 5], inhalation of radon was not formally included in the scope of the investigation. However, the results of separate calculations for airborne radon concentrations are given.
1.4. Structure
Following this introductory section, Section 2 provides a summary of the empirical evidence collected for this study pertaining to the characteristics of the various NORM residue deposits and their potential leachate characteristics. Section 3 provides details of the models and data used to perform the exposure pathway assessment and presents the results of the calculations.
Section 4 discusses the results of the dose assessment and draws some conclusions on the key exposure pathways and radionuclides involved. Details of the dose calculations, the relevant air dispersion considerations and the sensitivity of the results to the soil–groundwater distribution coefficients are given in three appendices.
2. EMPIRICAL EVIDENCE 2.1. Types of residue considered
Many bulk NORM residues contain elevated levels of natural radionuclides (see Table 1, for instance). Such materials include waste rock from many different types of mining operations and mineral processing residues such as ‘red mud’ from bauxite production, gold tailings, copper slag, tin slag and phosphogypsum (a by-product of fertilizer production). Uranium and thorium series radionuclides are commonly found in a number of mining operations including amongst others, uranium mining (e.g. Canada, Germany), mining of rare earth and niobium deposits (many sites internationally), gold mining (e.g. South Africa), tin mining (e.g.
Canada, Brazil) and copper mining (e.g. Australia). Mining of other non-metallic ores, along with ore processing, also generates large quantities of NORM, examples being the production of phosphate fertilizer and abrasives and the generation of fly ash and coal ash residues from coal-fired power generation.
The chemical and mineralogical properties of mining residues and metallurgical products and the presence of other chemicals such as organic matter or iron oxide can affect the availability of the radionuclides to the environment. Also, mining or metallurgical processes can affect the mobility of radionuclides in the natural environment. In this regard, there are few differences (other than differences in radionuclide activity concentration) between residues from the mining and processing of uranium ore and those from other types of mining and mineral processing operations. The resulting radiological impacts per unit activity concentration are likely to be similar for both cases.
Residues from uranium mining include waste rock, tailings, treatment sludges and so-called
‘clean waste’. Clean waste has no specific definition but based on precedent, as for example in Canada, would include materials such as waste rock with less than 0.03% U3O8 (implying a
238U concentration of about 3 Bq/g) and would be non-acid-generating. Other elements such as arsenic, selenium, sulphur (as sulphide), nickel, molybdenum and vanadium may also be present and may be used as a marker to define clean waste. Clean waste from uranium mining is likely to be similar to NORM residues with low mineralization that may be generated at other mining operations.
Potential issues with waste rock include elevated levels of radioactivity (radon, elevated gamma fields, and uranium/radium in leachates). In some cases, such materials also have the potential to generate acid drainage as a consequence of the oxidation of sulphide mineral present in the rock. In most jurisdictions, waste rock is closely regulated using various measures including monitoring, closure management and restrictions on its use as a product.
There are several approaches to the management of residues with acid generating and/or metal leaching potential. Short term management can include temporary storage on lined pads with all seepages intercepted and treated. Long term management includes isolation under engineered covers and disposal in mined out cavities. In some locations where local conditions permit, water barriers are used to prevent oxidation of sulphides and to reduce metal mobility. There is a reasonably extensive database on leachate chemistry from uranium mine residue deposits, but considerably less information on the quality of leachates from other types of NORM residue deposit.
2.2. Sources of data
The information regarding the characteristics of NORM residue deposits was obtained from the following sources:
• IAEA publications;
• Open literature and journal papers;
• Proceedings of workshops and conferences;
• SENES in-house reports and information from uranium mining and milling operations.
2.3. Characteristics of residues
The characteristics of NORM residue deposits are quite variable and are dependent in large part on the geological setting of the source material, the specific uranium/thorium content of the material and the effects of processing (concentration/mobilization of radionuclides in some cases). For most unaltered residue materials, radionuclides are mobilized with weathering and the passage of water and air through the pile. Some of the radionuclides will migrate more readily than others which may then be precipitated or adsorbed on the surfaces of residue materials and local subsurface soils. Retardation of the movement of radionuclides such as radium and uranium has been demonstrated at many locations and retention coefficients, although highly variable, have been reported for many soils and rock types.
2.3.1. Physical and chemical characteristics
The physical, chemical, and geochemical characteristics of mining and industrial mineral based residues vary greatly. A large amount of information is available from various international studies on such materials. Examples of important properties include geochemical components, the presence of chemical contaminants, acid generating potential, carbonate content, pH, water content, porosity, organic matter content, clay content and cation exchange capacity.
Data from a waste rock pile at the Rabbit Lake uranium mine in northern Saskatchewan, Canada indicates that this rock is comprised primarily of sandstones (with a smaller fraction of chloritic gneiss and graphitic gneiss) and approximately 50% of the material in the pile is greater than 10 cm in diameter [6]. Since the coarse fractions will retain much less moisture than the fines, the volumetric moisture content is expected to be substantially lower than that of the fines. For this waste rock pile, the average gravimetric moisture content was 8.7%, similar to the volumetric moisture contents in the range of 8% for another site in Saskatchewan (Key Lake). Moisture data from other waste rock sites are typically in the range 7–10% [6].
Waste rock from a former uranium mining area in Ronneburg, Germany contains carbonates, ochre limestones, and calcareous slates. These constituents are lumped together with slates and alum slates containing pyrite and carbon creating geochemical variabilities in the waste rock piles [7].
The waste rock pile at Barzava uranium mine in western Romania contains grey sandstone;
micro-conglomerates; clay with organic substances and crystalline schistose, with a particle size range of 10–100 mm [8].
The residues in the vicinity of the Kalna abandoned uranium mines in Serbia are made up of varied lithological complexes that differ in age, origin, mineral and petrographic
compositions. The oldest rocks are amphibolite, gneiss and greenstone of Riphean/Cambrian age (diabase phyllite formation) [9].
A large variety of NORM residues arising from the processing of ores includes tailings, treatment sludges, various slags from the production of tin, niobium and tantalum ores, as well as phosphogypsum, a by-product of fertilizer production.
2.3.2. Radioactivity content
In addition to the uranium, thorium and radium isotopes included in the uranium and thorium decay series, other radionuclides of potential concern in NORM residue deposit seepage include 210Pb and 210Po. Monitoring of groundwater around tailings basins has demonstrated that 210Pb and 210Po tend to be present at concentrations below that of 226Ra. Typically, 210Po appears immobile as a result of precipitation or adsorption on solid surfaces.
2.3.2.1. Residues from the mining and processing of uranium ore
A range of uranium and thorium concentrations for waste rock at Key Lake uranium mine, Canada are shown in Table 2. Similar results for waste rock at Rabbit Lake uranium mine, Canada are shown in Table 3 [6].
TABLE 2. TYPICAL RADIOACTIVITY LEVELS IN VARIOUS FORMS OF WASTE ROCK AT KEY LAKE URANIUM MINE, CANADA
Activity concentration (Bq/g)
Low grade ore ‘Special waste’ ‘Clean waste’ ‘Acidic waste’
U-238 >12 3.7–12 <3.7 (mean 1.2) 1.2
Th-232 >1.2 0.41–1.2 0.12 0.12
Note: These data were reported in the January 2005 waste rock management plan for Key Lake uranium mine.
The data are similar to those for other uranium mine waste rock piles.
TABLE 3. TYPICAL RADIOACTIVITY LEVELS IN WASTE ROCK AT RABBIT LAKE URANIUM MINE, CANADA
Unit Concentration
Uranium % U 0.0033-0.14
U-238 Bq/g 0.41-17
Ra-226 Bq/g 0.6-15
In Ref. [10], it is reported that test samples from waste rock piles from Midnite mine (a former uranium mine in the USA) indicated that the uranium content of the waste rock material ranged from 7 to 229 ppm (238U: 0.09–2.8 Bq/g).
In Germany, at the former Gessenhalde uranium heap leaching facility, the average residual soil concentrations of uranium, thorium, and lead were about 8.5, 12 and 20 ppm, respectively (238U: 0.1 Bq/g; 232Th: 0.04 Bg/g) [11].
The analytical results from residues in the vicinity of the Kalna abandoned uranium mines in Serbia indicated that uranium concentrations ranged from 1.4 to 4.3 ppm while thorium
concentrations ranged from 3.4 to 12.3 ppm. The results of the analysis indicated that the primary source of uranium in the area is the fragmented granitic rocks [9].
The compositions of the waste rock pile and other residues in the close vicinity of the Barzava uranium mine located in the west of Romania were measured. The analytical results indicated that the uranium concentration ranged from 53 to 175 ppm (238U: 0.65–2.2 Bq/g) while 226Ra concentrations were between 0.15 and 0.9 Bq/g [8].
Overall, a review of the information shows that the uranium activity concentrations in residues associated with the mining and processing of uranium ore generally vary between 0.1 and 17 Bq/g while the corresponding range for thorium is between 0.04 and 1.2 Bq/g. Thus, the uranium activity concentrations vary by more than two orders of magnitude, while the range of thorium concentrations is narrower. In both cases, the ranges extend above and below the regulatory criterion of 1 Bq/g (see Section 1.1).
2.3.2.2. Residues from the mining and processing of ores other than uranium ore
Significant concentrations of uranium and/or thorium are commonly associated with a number of other mining operations, including rare earth and niobium mining (many sites internationally), tin mining (several locations), copper mining (e.g. Olympic Dam, Australia) and gold mining (e.g. South Africa). Depending on the nature of the residues, the facilities containing the residues may or may not have engineered containment. Typical data on selected tailings and waste rock sites with NORM are provided in Table 4.
TABLE 4. TYPICAL RADIOACTIVITY LEVELS IN NORM RESIDUES OTHER THAN THOSE ASSOCIATED WITH THE MINING AND PROCESSING OF URANIUM ORE
Activity concentration (Bq/g) Waste rock,
copper mining
Phospho- gypsum [12]
Gold tailings, South Africa [13]
Red mud [14]
Coal ash
Waste rock, niobium mining
U-238 1.0
(0.11–2.1)
~ 0.1 (0.005–0.5)
1.0 (0.1–5)
<0.4 0.10–1.0 0.34
Th-232 – 0.004–0.6a – 0.41 – 1.2
Ra-226 – ~1
(0.015–5.1)
1.4 (0.1–5)
<0.18 – –
a Highly dependent on the deposit characteristics, with phosphogypsum from igneous sources containing much more thorium than phosphogypsum from sedimentary sources.
During the phosphoric acid production process, uranium, thorium and lead distribute primarily in the phosphoric acid, while most of radium, polonium, and traces of uranium appear in the phosphogypsum [15]. An analysis of 238U, 230Th and 226Ra in samples taken from three phosphogypsum stacks in Florida, USA and two in Canada is reported in Ref. [16].
The 238U activity concentrations were in the range of 0.092–0.530 Bq/g, while the activity concentrations for 230Th were in the range 0.072–0.150 Bq/g. The range of activity concentrations for 226Ra was 0.310–0.930 Bq/g. These data confirm that some preferential distribution to the fertilizer product of uranium occurs compared with radium, most of which ends up in the phosphogypsum by-product.
An analysis of phosphogypsum derived from phosphate rock from Florida Togo, and Idaho indicated that the average activity concentrations of 226Ra ranged from 0.43 to 0.93 Bq/g, while the average activity concentrations of 210Pb ranged from 0.34 to 0.840 Bq/g [17]. The average activity concentrations of 228Th were between 0.004 and 0.009 Bq/g.
For NORM residues other than those associated with the mining and processing of uranium ore, the range of uranium activity concentrations is generally from 0.1 to 5 Bq/g, while the corresponding range for thorium activity concentrations is 0.004 to 1.2 Bq/g. The activity concentrations depend on the type of mining and processing activity and type of residue. As with uranium mining residues, the ranges of concentrations extend above and below the recommended regulatory criterion of 1 Bq/g.
2.3.3. Leachate
There is often a potential to produce contaminated drainage even though the residue piles are believed to contain only low levels of contaminants. A thorough understanding of the chemical and mineralogical characteristics of the residue is necessary to understand and predict leachate chemistry.
NORM residues from mineral processing operations typically contain meta-stable components such as precipitates that can release contaminants including heavy metals and radionuclides. Also, chemical oxidation can have a large impact on the mobility of radionuclides in tailings and other residues. Uranium, for example, may be converted from U(IV) to U(VI) and become more mobile especially in the presence of carbonate. A fraction of the radium content may be mobilized in hydrometallurgical processes and co-precipitated with other metal sulphates. The mobility of this co-precipitated radium becomes controlled by sulphate levels. Dissolution is assumed to be the dominant mechanism of release of uranium and 226Ra from oxide, sulphate and carbonate minerals. However, release from these minerals can also be influenced by the pH and redox changes associated with oxidation. The presence of carbonate, iron oxides, clay minerals and organic matter can greatly affect the mobilization of radionuclides in the residues. Because of the stability of the uranyl carbonate complexes, uranium-rich residues with naturally high carbonate contents tend to produce higher levels of uranium in leachates. High levels of chloride also tend to mobilize radium.
Fluctuations in pH can result in a relatively rapid release of radionuclides and this is likely to be attributable to changes in both the solubility of the host mineral and the retention coefficients. For example, uranium and thorium are more soluble under acid conditions while radium solubility may be reduced (in the presence of high sulphate concentrations). Should a residue deposit be acid generating, uranium levels may increase, thorium will be mobilized if the residue deposit is very acidic, and radium levels will be at similar or possibly lower levels owing to secondary precipitation of radium–sulphate complexes.
Overall, the factors controlling the solubility of key radionuclides in NORM residues are reasonably well known but the actual characteristics of leachate vary greatly. Laboratory and field data from various sites show that the rates of contaminant release from waste rock change over time. In cases where there is no sharp change in pH, that is, where the system remains neutral, contaminant release rates generally decrease over the long term [6].
In addition to the above factors, water flow within NORM residue deposits is another important factor influencing the mobility of radionuclides from these deposits. The average
velocity at which water moves through the deposit is a function of various parameters, including the infiltration rate and the overall volumetric water content of the deposit.
Typical leachate characteristics for residues at the Key Lake uranium mine in Canada are shown in Table 5. These residues produce neutral drainage with elevated levels of uranium isotopes and 226Ra and negligible levels of thorium. Uranium levels show the greatest variability.
TABLE 5. TYPICAL CHARACTERISTICS OF LEACHATES FROM URANIUM MINING Leachate from ‘clean waste’ Leachate from ‘acidic waste’
U (mg/L) 0.5 (0.1–5) 5
U-238 (Bq/L) 6 (1.2–60) 60
Ra-226 (Bq/L) 0.5 (0.1–1.5) 0.5
Th (µg/L) <1 100
Th-232 (Bq/L) <0.004 0.4
Note: The values are based on data reported in the January 2005 waste rock management plan for Key Lake uranium mine. These data are similar to data for other mines.
The leachate from waste rock piles from the Midnite mine in the USA exhibited alkaline characteristics with leachate pH values near to 8.5 [10]. The tests indicated that the oxidation during the dry period increases the uranium leaching during subsequent rainfalls. The uranium leach concentrations were mostly less than 5 mg/L (238U: 60 Bq/L) and generally about 2 mg/L (238U: 25 Bq/L). On the other hand, the samples from different layers of the waste rock pile had pH ranges of 3 to 7. The uranium concentrations of leachate were consistently below 0.2 mg/L (238U: 2.5 Bq/L) with initial leachate concentrations as high as 2.9 mg/L (238U: 36 Bq/L) [10]. This part of the waste rock pile is not as reactive with respect to uranium.
The test results for leach samples from the Rabbit Lake waste rock pile in Canada indicate that arsenic, molybdenum, nickel and uranium dominate the chemistry in several of the samples [6]. The alkalinity was typically very low, and sulphate was the dominant anion. The pH values tended to be weakly acidic. Table 6 shows the test results for leach concentrations in samples of Rabbit Lake waste rock.
TABLE 6. TYPICAL CHARACTERISTICS OF LEACHATE FROM RABBIT LAKE URANIUM MINE [6]
Unit Concentration
U µg/L 40–13 800
U-238 Bq/L 0.5–170
Pb mg/L <0.002–0.022
Ra-226 Bq/L 0.33–5.5
Lead that is leached from NORM residue deposits may adsorb on underlying soils. Lead may precipitate in soils if soluble concentrations exceed about 4 mg/L at pH 4 and about 0.2 mg/L at pH 8. In the presence of phosphate and chloride, these solubility limits may be as low as 0.3 mg/L at pH 4 and 0.001 mg/L at pH 8 [18]. Therefore, in situations in which concentrations of lead exceed these values, the estimated distribution coefficients (Kd) may reflect precipitation reactions rather than adsorption reactions. Anionic constituents such as phosphate, chloride, and carbonate are known to influence lead reactions in soils either by precipitation of minerals of limited solubility or by reducing adsorption through complex formation [18]. Depending upon the nature of the tailings or other residues, the deposits may or may not have engineered containment.
It is also reported in Ref. [7] that the analytical prediction of the geochemical behaviour of secondary minerals at the Ronneburg site is very complex and quantitative predictions of the water quality in the seepage are notoriously unreliable.
Typical leachate data for selected tailings and waste rock sites are given in Table 7. The analytical data for mine and seepage water from the Wismut mining district located in Ronneburg, Germany are shown in Table 8.
TABLE 7. TYPICAL LEACHATE ACTIVITY CONCENTRATIONS ASSOCIATED WITH NORM RESIDUES
Activity concentration (Bq/L) Waste rock,
copper mining
Phospho-
gypsum Red mud Coal ash Waste rock
niobium mining
U-238 2.4 0.043a 0.1–0.6 – <0.2
Ra-226 0.26 <0.74a
0.07–0.53b
0.02–0.03 0.8 1
a From Ref. [15].
b From Ref. [19].
TABLE 8. COMPOSITION OF TYPICAL MINE AND SEEPAGE WATERS OF THE RONNEBURG SITE, GERMANY [7]
pH
Uranium concentration
(mg/L)
Activity concentration
(Bq/L) U-238 Ra-226 Mine water, central section, measuring point e-567 3.8 0.1 1.2 0.150 Mine water, central section, influenced by open pit
mine and dumps, measuring point e-480
2.9 4.4 55 0.384
Mine water, SE section of deposit, measuring point MW 435/2
7 1.7 21 0.192
Seepage, Absetzerhalde dump, measuring point e-440 2.8 7.2 89 <0.01 Seepage, Beerwalde dump site, measuring point s-611 7.6 5.2 64 0.136
The concentrations of radionuclides in the Lerchenbach creek north of the former tailings management area of a Wismut site in East Trunzig, Germany are shown in Table 9 [20]. The tailings were generated from ore containing 238U decay series radionuclides at an activity concentration of the order of 10 Bq/g. It is concluded that, given the characteristics of waters encountered in the vicinity of abandoned Wismut sites (with mostly neutral or acidic pH), the radionuclide transport was typically determined by the uranium isotopes 238U, 234U and 235U.
These radionuclides accounted for the major part of the effective dose.
TABLE 9. ACTIVITY CONCENTRATIONS IN SURFACE WATER NEAR THE TAILINGS MANAGEMENT AREA, EAST TRUNZIG, GERMANY [20]
Activity concentration in water (Bq/L)
U-238 5.2
U-234 6.1
Th-230 0.17
Ra-226 0.02
Pb-210 0.025
Po-210 0.025
U-235 0.24
Pa-231 0.015
Ac-227 0.015
It is reported in Ref. [8] that the uranium concentration in surface water in the close vicinity of Barzava uranium mine in Romania was 0.014 mg/L (238U: 0.17 Bq/L), while the activity concentration of radium was 0.043 Bq/L. The authors concluded that the argillaceous soil at the bottom of the waste rock pile worked as a barrier that retarded the movement of radioactive contaminants toward the ground and consequently surface water.
The analytical results for well water in the vicinity of the Kalna abandoned uranium mines in Serbia are reported in Table 10 [9]. It was concluded that the geochemical barriers of clays and organic materials (alluvium) and faster filtration of groundwater control the dissolution of uranium and trace elements in groundwater downstream of the mines. Uranium concentrations were all less than 0.009 mg/L (0.1 Bq/L). A radiometric analysis of samples from the mine residue was not provided in this study.
TABLE 10. RANGES OF MEASURED ELEMENTS IN WELL WATER IN VARIOUS LOCATIONS AROUND THE KALNA ABANDONED URANIUM MINE
Uranium concentration
(mg/L)
Activity concentration (Bq/L)
238 pH
U 226Ra
Balta Berilovac 0.005–0.009 0.062–0.111 0.05–0.07 7.3–7.4
Vrtovci 0.001–0.002 0.012–0.025 0.09–0.20 7.0–7.1
Inovo 0.002–0.008 0.025–0.100 0.05–0.06 7.4–7.7
Gornja Kamenica 0.001–0.002 0.012–0.025 0.20–0.60 7.3–7.5
Donja Kamenica 0.001–0.003 0.012–0.037 0.10–0.19 7.5–7.8
Strbac 0.001–0.003 0.012–0.037 0.05–0.07 7.1–7.3
Baranica 0.001–0.003 0.012–0.037 0.11–0.20 7.4–7.6
During the uranium milling periods at Moab uranium mine site in Utah, USA, Atlas Minerals monitored 226Ra, 230Th, 222Rn and uranium at the tailings pond [21]. The results indicated that the 226Ra and 230Th activity concentrations were 3.7 and 1.9 Bq/L, respectively. The uranium activity concentration was 22 Bq/L.
Data published in Ref. [22] for typical seepage/groundwater levels measured in the plume from the Nordic tailings management area (Elliot Lake, Ontario, Canada) are shown in Table 11. The average radionuclide activity concentration of the Nordic tailings is about 10 Bq/g [23].
TABLE 11. ACTIVITY CONCENTRATIONS IN SEEPAGE/GROUNDWATER AT VARIOUS DISTANCES FROM THE NORDIC TAILINGS DAM, ELLIOT LAKE, CANADA
Distance from
dam (m) pH Activity concentration (Bq/L)
226Ra 238U 210Pb 232Th 230Th
0 4.5 5.5 6.5 4 Not detectable 0.14
10 4.5 2.0 1.2 Not detectable Not detectable 0.14
20 5.0 0.4 <0.1 Not detectable Not detectable Not detectable 70 6.3 <0.1 <0.1 Not detectable Not detectable Not detectable A mathematical modelling of leaching of radionuclides from the Urgeirica uranium tailings located in central Portugal is reported in Ref. [24]. The modelling results indicated that the activities of 226Ra and total U in well water 500 m away could reach 1 and 6 Bq/L, respectively, after 100 years of leaching and movement of these radionuclides through the groundwater. Based on the modelling results, it was concluded that due to the slow rates of contamination migration, only radionuclides with relatively long half-lives are important in the transport process. Therefore, with regard to polonium, a nil concentration was estimated in the well water. This can be explained by the fact that the half-life of 210Po is about 137 d, which is very short compared with its travel time. The radionuclides 230Th and 210Pb are generally not transported over significant distances due to the particle-reactive nature of thorium and the great tendency of lead to be adsorbed by the aquifer sediments [24].
In Germany, at the former Gessenhalde uranium heap leaching facility, the average leachate concentrations of uranium and lead were 59 and 0.9 µg/L, (0.73 and 0.11 Bq/L), respectively [11]. It was concluded that, based on the current level of contamination and the available historical information, leachate infiltrated through the barrier soil into the loamy soil and was trapped on top of the glacial clay, thereby being prevented from infiltrating further downwards. The water concentrations were found to be pH dependent for most elements, with low pH areas exhibiting higher concentrations. Most elements indicated an almost logarithmic increase with decreasing pH [11].
Refs [25], [26] describe a geochemical and radiological assessment of a uranium tailings dam in Poços de Caldas, Brazil. The activity concentrations of 238U series radionuclides in the solid material were in the range 4.1–6.6 Bq/g. Radionuclide concentrations in near-surface seepage and in the tailings dam pond water are summarized in Table 12. Although the data are limited, they show that the pond water values are very much lower than the seepage values.
For example, the 226Ra activity concentrations differ by a factor of about 40, while 210Pb concentrations differ by a factor of 200. It was also noted that the activity concentrations of
238U and 226Ra in the groundwater were essentially the same upstream and downstream of the point at which the effluents from the tailings dam were released because the effluents were treated with lime and BaCl2.
Ref. [27] gives a compilation of the results of studies conducted over several years to determine radionuclide concentrations in waters impacted by gold mining operations in the sedimentary gold deposits of the Witwatersrand Basin in South Africa. The area is characterized by low grade uranium mineralization with an average uranium oxide grade from a particular conglomerate formation in a mine varying from 0.001 to 0.078 %. The mean activity concentrations of 238U and 226Ra in the mine tailings were each about 1 Bq/g. The mean activity concentrations of 238U and 226Ra in surface water and groundwater samples are given in Table 13. The activity concentrations of measured radionuclides other than uranium isotopes and 226Ra were close to natural background concentrations.
Heavy metals and other toxic elements and radionuclides are not readily leachable from phosphogypsum solids [16]. It was noted, however, that process fluids contained in the pore spaces of the phosphogypsum stacks can be a source of groundwater contamination.
TABLE 12. ACTIVITY CONCENTRATIONS IN SEEPAGE WATER AND POND WATER AT A URANIUM MINE TAILINGS DAM IN BRAZIL
Activity concentration (Bq/L)
Seepage water (n=20) Pond water (n=1)
Ra-226 1.27 (0.33–5.0) 0.03
Ra-228 <1.30 0.38
Th-232 0.10 (0.01–0.7) <0.015
U-238 2.89 (0.15–11) 0.22
Pb-210 5.3 (0.07–38) <0.02
TABLE 13. ACTIVITY CONCENTRATIONS IN SURFACE WATER AND
GROUNDWATER CONTAMINATED BY GOLD TAILINGS IN SOUTH AFRICA [27]
Mean activity concentration (Bq/L)
Groundwater Surface water
At edges of tailings dams
Surrounding aquifers
Mine discharge
Tailings return
Surrounding streams, lakes and
dams
U-238 1.22 0.10 1.19 3.44 0.37
Ra-226 0.04 0.02 0.29 0.72 0.04
It is reported in Ref. [15] that groundwater samples taken from the vicinity of phosphogypsum stacks in Florida, USA exhibited activity concentrations of 0.002–
0.043 Bq/L for 238U, 0.002–0.021 Bq/L for 230Th and 0.004–0.74 Bq/L for 226Ra. It was noted that the phosphogypsum slurry is discharged into the stacks in an acidic state and that the characteristics of the pore water in the phosphogypsum stacks are determined by the phosphoric acid production process rather than the derived leachate from solids.
A 226Ra activity concentration of 0.07–0.53 Bq/L was measured in leachates from phosphogypsum stockpiles and it was found that the leaching of radium may be slow in field conditions near the stockpiles [19]. However, considering the large quantity of phosphogypsum disposed of at individual sites, contamination of surface water and groundwater was considered to be of concern at that locality. In another study, the leachates from Togo phosphate rock were found to have 226Ra concentrations of 0.23–0.55 Bq/L [28].
Few data are available on the concentrations of 210Pb and 210Po in leachate from NORM residue deposits. However, data from the monitoring of seepage at uranium tailings basins may be used as an indication of the potential levels of these radionuclides compared with radium. For neutral seepages, concentrations of 210Pb and 210Po tend to be <20% and <5% of the 226Ra concentrations, respectively. In acidic seepages, concentrations of 210Pb are similar to those of 226Ra, while 210Po concentrations are about 10%.1 This will vary greatly by the type of NORM residue but nonetheless the uranium mining data provide an indication of what levels might be expected in NORM residues.
A study of contamination in a subsurface disposal area (SDA) at the Idaho National Engineering Laboratory (INEL) in south-eastern Idaho is reported in Ref. [29]. The SDA had been used since 1952 to dispose of radioactive wastes, some of which contained 238U, 226Ra and 210Po. In this study, multi-media sampling and analysis indicated that the 210Po levels detected in all media collected within and adjacent to the SDA were not statistically different from those concentrations detected in control area samples. The low concentrations of 210Po detected in SDA media indicate that the disposal of radioactive wastes has not resulted in elevated 210Po levels in water, surface soil, vegetation or small mammal tissues in the surrounding environment. The highest 210Po concentration in soils (0.23 ± 0.06 Bq/g) was statistically greater than all other sampling locations (range of means was 0.03–0.1 Bq/g).
The data reviewed above suggest that although there is considerable variability from one residue deposit to another, the leachates are of roughly similar quality.
For uranium mining residues, this review suggests that the uranium activity concentrations of most leachates are within the range 0.5–170 Bq/L while the thorium activity concentrations are typically very much lower, of the order of 1.9 Bq/L or less. The activity concentrations in groundwater are generally within the range 0.005–55 Bq/L. Comparing the range of uranium concentrations in uranium mining residues summarized in Section 2.3.2 (0.1–17 Bq/g) with the corresponding range of leachate activity concentrations discussed above (0.5–170 Bq/L), it can be concluded that the range of leachate concentrations (in Bq/L) is roughly an order of magnitude higher than the range of uranium activity concentrations (in Bq/g) in the residues concerned. The higher leachate activity concentration of 170 Bq/L is for acidic waste rock leachate in Rabbit Lake uranium mine.
For non-uranium mining residues, the uranium activity concentrations of the leachate are generally within the range 0.1–2.4 Bq/L. Comparing this range of values with the range of 0.1–11 Bq/g in residues, it can be concluded that for every becquerel per gram of uranium activity concentration in the residue, the leachate concentration is less than 1 Bq/L.
For thorium, for every becquerel per gram of activity concentration in the residue, the leachate activity concentration in the leachate is generally less than 1 Bq/L.
1 Data obtained from groundwater monitoring data at Denison and Stanrock uranium mines, Elliot Lake, Canada.
2.3.4. Radionuclide transport in groundwater
The transport of radionuclides in groundwater has been studied extensively. Almost without exception, rocks and soils attenuate radionuclides when the concentrations of such radionuclides in the source exceed background levels.
The transport process of soluble materials in water generally has two components: advection and dispersion. Advection is transport along with the average pore water velocity. Dispersion is transport within the water, due to both molecular diffusion and small scale differences in flow speeds. When the pore water velocity is high, the dispersion component is not a significant contributor to the movement of the soluble chemicals. This is particularly true when the groundwater is moving in coarse sandy soil, where the pore water movement is relatively fast, and the retardation factor is relatively small. A range of hydraulic conductivities and permeabilities for various rocks and unconsolidated deposits can be found in Ref. [30]. The hydraulic conductivity K ranges from 0.1 m/s for coarse gravel to 10–13 m/s for metamorphic and igneous rock. It is expected that the hydraulic conductivity for residue materials with large particle sizes would have a wide range depending on profiles and surface conditions of the residue deposits; the nature, size range and size segregation of the materials;
the pore volumes; compaction; and climatic conditions such as freeze–thawing, wetting and drying.
The retardation factor (a function of the soil–water distribution coefficient, Kd) for natural materials can vary over orders of magnitude and is dependent on numerous factors including pH, grain size, cation exchange capacity (CEC), redox conditions, salinity and other groundwater chemical characteristics. Without site-specific testing, applying a distribution coefficient in a contaminant transport model is subject to uncertainty.
The strong effect of the distribution coefficient, which reflects the partitioning of the radionuclide concentration between the solid and liquid phases, was demonstrated by a study at the site of the former White King uranium mine in the USA [31]. Although no actual distribution coefficients were calculated at the White King site, no down-gradient uranium was detected, even though pore water uranium concentrations in the stockpile were about 1000 Bq/L. The overburden stockpile pore water concentrations were less than 0.67 Bq/L, with a concentration of only 2.87 Bq/L immediately under the stockpile. Thus, with an inferred high distribution coefficient at this site, the uranium appears to be quite immobile. No radium was detected in the leachate, apparently having been contained within metal sulphates.
One example of groundwater migration from a uranium tailings deposit is the Nordic tailings deposit in Elliot Lake, Canada, where it was found that the migration of uranium and radium from acid leachates was greatly retarded in the groundwater aquifer below the deposit [22]. In the inner acid plume, radium and uranium distribution coefficients were calculated to be 70 and 2000 mL/g, respectively. It is likely that the migrating acid plume is being neutralized by alkalinity in soils. In the neutral outer zone of the plume, distribution coefficients for uranium were reported to be about 3600 mL/g.
Leaching rates for residue deposits imply that typically only a very small fraction of the inventory of NORM constituents is mobilized. As such, leaching is expected to continue for long time periods and retardation is likely to only provide a delay in the time at which the peak contaminant levels will reach the receptor. Therefore, for modelling purposes, the peak long term source concentrations in the absence of radioactive decay are likely to be similar to the concentration in the diluted source. This would not be the case if the radionuclides were
permanently removed (that is, if the reactions were not reversible). This can occur with elements such as uranium which under reducing conditions may change its oxidation state and hence reduce the solubility (availability) of the mineral.
2.3.5. Distribution coefficients
In an experimental study to evaluate distribution coefficients for a soil–groundwater system, values were obtained for five types of soils that resemble NORM residue deposit materials [32]. Table 14 gives these values for U and Th.
At concentration ranges that occur in the leachate from residue deposits, the extent of thorium adsorption can be estimated from the soil pH. Studies have shown that the lowest thorium distribution coefficient was 20 mL/g for a measurement made on a pH 10 soil, while the largest thorium distribution coefficient was 170 000 mL/g for a measurement made on a silt–
quartz soil of schist origin [18]. Based on the studies reviewed in Ref. [18], thorium distribution coefficients were estimated for various pH ranges and are given in Table 15.
TABLE 14. DISTRIBUTION COEFFICIENTS FOR U AND Th IN VARIOUS SOIL TYPES [32]
Distribution coefficient (mL/g)
U Th
Glacial till: Sandy silty clay with included fine to coarse gravel with very occasional cobbles.
46 24 000
Sand: Fine to medium sand 560 280
C.1.2: Slightly silty fine to coarse sand and mainly fine with occasional cobbles
46 5800
C.3: Medium to coarse sand with a lot of fines with occasional cobbles
900 280
C.6: Largely coarse grained sand with a lot of fines with numerous cobbles and small boulders
2200 5800
TABLE 15. RANGES OF DISTRIBUTION COEFFICIENTS FOR THORIUM [18]
Distribution coefficient (mL/g)
pH 3–5 pH 5–8 pH 8–10
Minimum 62 1700 20
Maximum 6200 170 000 2000
A compilation of many studies on uranium partitioning indicates that pH and dissolved carbonate concentrations are the two most important factors influencing the adsorption behaviour of U(VI), the most mobile species of uranium [18]. Table 16 shows the range of estimated minimum and maximum distribution coefficients for uranium for partitioning between water and soil and crushed rock.
TABLE 16. RANGES OF DISTRIBUTION COEFFICIENTS FOR URANIUM [18]
Distribution coefficient (mL/g)
pH 3 pH 4 pH 5 pH 6 pH 7 pH 8 pH pH 10
Minimum <1 0.4 0.25 100 63 0.4 <1 <1
Maximum 32 5000 160 000 1 000 000 630 000 250 000 7900 5 The indications in Ref. [18] are that the lower bound of the range of uranium distribution coefficients is based on values estimated for quartz. It is unlikely that actual distribution coefficients for U(VI) can be much lower than those represented by this lower bound.
Ref. [18] also cites field-derived uranium distribution coefficients for 238U and 235U from Ref. [33] and plots the derived values versus pH. As shown in Fig. 1, the uranium distribution coefficients vary from 1.2 to 34 000 mL/g over a pH range of approximately 3–6.7.
A summary of distribution coefficients for lead adsorption on soils from various studies is provided in Ref. [18] and is shown in Table 17. This information confirms that within the pH range of soils (4–11), lead adsorption increases (as does precipitation) with increasing pH.
In support of the US Department of Energy RESRAD modelling system for estimating doses, a great deal of information has been compiled, including information on distribution coefficients (see, for instance, Ref. [34]). As discussed in Section 3.1.2, distribution coefficients from the RESRAD data collection are compared with values based on the leachate and residue concentrations cited earlier in this Section to provide a basis for distribution coefficients that are considered to be reasonably representative of a typical NORM residue deposit.
FIG. 1. Field-derived distribution coefficients plotted as a function of pore water pH for contaminated soil and pore water samples. Square and circle symbols represent field-derived distribution coefficients for 238U and 235U, respectively. Solid square and circle symbols represent minimum distribution coefficients for 238U and 235U, respectively, that were based on minimum detection limit values for the concentrations for the respective uranium isotopes in pore waters associated with the soil sample. (From Ref. [18])
